Role of Dust and Iron Solubility in Sulfate Formation during the Long-Range Transport in East Asia Evidenced by 17O-Excess Signatures

Numerical models have been developed to elucidate air pollution caused by sulfate aerosols (SO42–). However, typical models generally underestimate SO42–, and oxidation processes have not been validated. This study improves the modeling of SO42– formation processes using the mass-independent oxygen isotopic composition [17O-excess; Δ17O(SO42–)], which reflects pathways from sulfur dioxide (SO2) to SO42–, at the background site in Japan throughout 2015. The standard setting in the Community Multiscale Air Quality (CMAQ) model captured SO42– concentration, whereas Δ17O(SO42–) was underestimated, suggesting that oxidation processes were not correctly represented. The dust inline calculation improved Δ17O(SO42–) because dust-derived increases in cloud-water pH promoted acidity-driven SO42– production, but Δ17O(SO42–) was still overestimated during winter as a result. Increasing solubilities of the transition-metal ions, such as iron, which are a highly uncertain modeling parameter, decreased the overestimated Δ17O(SO42–) in winter. Thus, dust and high metal solubility are essential factors for SO42– formation in the region downstream of China. It was estimated that the remaining mismatch of Δ17O(SO42–) between the observation and model can be explained by the proposed SO42– formation mechanisms in Chinese pollution. These accurately modeled SO42– formation mechanisms validated by Δ17O(SO42–) will contribute to emission regulation strategies required for better air quality and precise climate change predictions over East Asia.


INTRODUCTION
The formation of sulfate aerosols (SO 4 2− ) in East Asia is a concern because it causes severe haze and pollution events, particularly in China, resulting in low visibility 1 and public health problems. 2 SO 4 2− is also one of the important shortlived climate forcers (SLCFs) related to regional and global climate changes. 3,4 Atmospheric SO 4 2− is produced mainly via the oxidation of sulfur dioxide (SO 2 ), 5 and SO 2 emissions from China accounted for approximately a quarter of global total emissions. 6 The problem of SO 4 2− pollution in East Asia is not confined to the emission sources in China because of the relatively long lifetime of SO 4 2− and the westerlies in the midlatitudes. 7 Thus, the long-range transport of SO 4 2− from the source region on the Asian continent to the downstream region of the air masses (i.e., Republic of Korea and Japan) has been thoroughly investigated. 8−10 To elucidate pollution caused by atmospheric SO 4 2− , chemical transport models (CTMs) have been developed and used in various studies. In CTMs, the gas-phase oxidation by the hydroxyl (OH) radical and the aqueous-phase oxidations by hydrogen peroxide (H 2 O 2 ), ozone (O 3 ), and O 2 catalyzed by transition-metal ions (TMIs) for SO 4 2− formation are typically considered. However, typical CTMs often underestimate the burden of atmospheric SO 4 2− , particularly in China, 11,12 suggesting that some SO 4 2− formation in the atmosphere is missing. To date, heterogeneous SO 4 2− production mechanisms particularly for aerosol surfaces 13−16 have been proposed to explain this missing formation such as the enhanced role of the oxidation by reactive nitrogen, 17 25,26 Nonetheless, the observational evidence has not yet identified a specific mechanism. For example, while the importance of NO 2 for SO 4 2− formation within the aerosol surface was proposed, 17 follow-up studies have cast doubt on the impact of its reaction. 25,26 Thus, these proposed reactions have been highly controversial arguments. These pollution and haze events over China and their impact on the downstream region should be mitigated, mainly via the reduction of anthropogenic SO 2 emissions. However, according to reports in Western countries, atmospheric SO 4 2− has declined less rapidly than would be expected from decreases in SO 2 emissions. 27,28 This unknown response has been attributed to chemical feedback mechanisms of a weakening H 2 O 2 limitation on the S(IV) + H 2 O 2 pathway 29 and acidity-driven enhancement of the S(IV) + O 3 pathway under low SO 2 conditions 30,31 in Western countries. Therefore, accurate CTMs for both concentration and oxidation processes are required to establish an effective emission regulation strategy for improving air quality, and the accurate implementations of atmospheric SO 4 2− formation mechanisms in CTMs are essential for elucidating present pollution and predicting future air quality. Yet, the atmospheric SO 4 2− formation pathways implemented in CTMs have been simply evaluated by total SO 4 2− masses and not validated by independent observational evidence.
One proven method to validate atmospheric SO 4 2− formation is the mass-independent oxygen isotopic composition (Δ 17 O) 32 38 Without the validation of SO 4 2− formation processes in East Asia, studies using current typical CTMs may not adequately predict air quality and climate change, given the possible cause of chemical feedback mechanisms, which have been studied in Western countries. 27 20,39,40 which is a unique aspect of the East Asian environment, is investigated because dust-driven higher pH alters acidity-dependent aqueous-phase SO 4 2− formation, such as TMI-catalyzed oxidation by O 2 and oxidation by O 3 . In addition, the impact of TMIs on SO 4 2− formation is examined because these are uncertain parameters for emissions and solubilities in the model. 41 Through comparisons of observed and modeled SO 4 2− and Δ 17 O(SO 4 2− ) at NOTOGRO in Japan, this study clarifies the key factors controlling the formation process of long-range-transported SO 4 2− in East Asia.

Observation of Aerosol and Oxygen Isotopic
Composition in Japan. The atmospheric observations were performed at NOTOGRO at 37.45°N, 137.36°E ( Figure 1). The Noto peninsula extends from the west coast of mainland Japan approximately 150 km into the sea of Japan, and NOTOGRO is located on the tip of this peninsula. The geographical location of NOTOGRO is ideal for capturing the atmospheric variation in East Asia because it is surrounded by the sea and isolated from major pollution sources in Japan. 42 The aerosol samples were collected by a high-volume air sampler (MODEL-120 SL, Kimoto Co., Ltd., Japan) mounted on the rooftop (∼10 m above sea level) of NOTOGRO. Fine (<2.5 μm) and coarse (>2.5 μm) samples were collected on prebaked (450°C for 4 h) quartz filters (2500QAT-UP, Pall Co., Ltd.; TE-230-QZ, Tisch Environmental Inc.). Sampling was performed at a flow rate of ∼1.05 m 3 /min, and the sampling interval was usually 1 or 2 weeks. After sampling, the filters were wrapped in aluminum foil, sealed in polyethylene bags, and stored in a clean freezer at −20°C prior to the measurement at the Tokyo Institute of Technology, Japan. In the laboratory, half of each filter was soaked in ultrapure water (30 mL) in a 50 mL centrifuge tube (Centricon plus-70, Millipore). The sample solution was separated from the insoluble materials and the filter by centrifuging in a centrifugal filter unit for 10 min. This method can recover >98% of the   31,37,43 Briefly, 1 or 2 μmol of H 2 SO 4 separated by ion chromatography was chemically converted to Na 2 SO 4 , and 30% of H 2 O 2 solution (1 mL) was added, and the mixture was dried. The Na 2 SO 4 was converted to silver sulfate (Ag 2 SO 4 ) using an ion-exchange resin. This Ag 2 SO 4 powder was transported in a custom-made quartz cup, which was dropped into the furnace of a high-temperature conversion elemental analyzer (TC/EA, Thermo Fisher Scientific) at 1000°C and thermally decomposed into O 2 . The O 2 gas was introduced separately into an isotope ratio mass spectrometer to measure m/z = 32, 33, and 34. The ) = 2.4‰) with the same procedure described previously. 31,37,43 In this correction for isotopic analysis, SD (1σ) for the corrected values for standard B was 0.11‰ based on measurements for the samples collected in 2015, and this 1σ uncertainty was used for the error of the isotopic measurements in the present study. The raw observation data are presented in the Supporting Information (Table S1). All data for SO 4 2− concentration and Δ 17 O-(SO 4 2− ) are corrected using Na + concentration to non-sea salt fraction of SO 4 2− (nss-SO 4 2− ) in a similar manner reported previously. 31 2.2. Regional Chemical Transport Modeling over East Asia. 2.2.1. Model Description. The regional air quality modeling was conducted with the CMAQ model version 5.3.1. 44,45 In this study, the simulation domain covered the entirety of East Asia with a horizontal resolution of 36 km ( Figure 1) and 44 nonuniform layers from the surface to 50 hPa. One gas-phase reaction and five aqueous-phase reactions in cloud are involved in SO 2 oxidation (i.e., SO 4 2− formation) in the original CMAQ. The one gas-phase reaction is SO 2 oxidation by OH (GAS), and the five oxidants in the aqueousphase reactions in cloud are H 2 O 2 (AQ(H 2 O 2 )), O 3 (AQ(O 3 )), O 2 catalyzed by TMIs (AQ(O 2 )), peroxyacetic acid (PAA) (AQ(PAA)), and methyl hydrogen peroxide (MHP) (AQ(MHP)). In addition to these five oxidants in the original CMAQ, taking into account the elevated NO 2 concentration in Chinese pollutions, 17−19 the aqueous-phase pathway in cloud via NO 2 was added in this study. This inclusion partly improved the model underestimation issue during winter in our previous study. 46 Note that the production of SO 4 2− on the aerosol surface is not considered in this study because the specific Chinese haze events that occurred in limited areas are out of scope and our focus is on capturing SO 4 2− pollution over East Asia from the viewpoint of the background site in Japan. This version 5.3.1 of CMAQ has several features to achieve the purpose of this study. At first, Fe and Mn emissions can be treated as independent variables, although these emissions had been previously dependent on the total PM 2.5 emission. Thus, we implemented the Transition Metal Inventory-Asia version 1.0 47 to calculate the emissions of Fe and Mn. Furthermore, although the default setting of CMAQ does not possess, we implemented the pH dependency rate constant for the calculation of AQ(O 2 ) using the synergistic relationship between Fe and Mn developed in our previous study. 46 These explicit treatments of Fe and Mn concentrations are major advantages of CMAQ and our study because other CTMs, such as GEOS-Chem, treat TMI concentrations as a fraction of the PM 2.5 concentration. Because of these improvements for Fe and Mn emissions and the rate constant for the calculation for AQ(O 2 ), we tested the remaining uncertainty of solubilities of Fe and Mn, as described in Section 2.2.2.
Second, it is possible to implement the developed physicsbased inline dust calculation after CMAQ version 5.2. Before this version, the effect of soil dust for neutralization and altering cloud drop pH could not be considered, even though one of the characteristics in East Asia is the role of mineral dust originating from the Taklamakan and Gobi Deserts, the Loess Plateau, and Inner Mongolia. However, dust is not implemented as the default setting of CMAQ, which hampers precise calculation for the pH-dependent SO 4 2− formation pathways, especially for AQ(O 3 ) and AQ(O 2 ). The impact on SO 4 2− formation caused by the difference with or without implementation of inline dust calculation was tested in this study, as explained in Section 2.2.2. The details of the modeling description and chemical configurations are, respectively, given in the Supporting Information (Sections S1 and S2).

Model Experiments.
In this study, we compared the following three experiments (Exp. A, Exp. B, and Exp. C) to investigate the role of mineral dust and solubilities of Fe and Mn for SO 4 2− formation, as summarized in Table 1. In Exp. A, which is the original settings in the CMAQ model version 5.3.1, the solubilities of anthropogenic Fe and Mn were set as 10 and 50% and dust inline calculation was not implemented (Table 1). In Exp. B, the modulation of pH by mineral dust was considered, and thus the newly developed dust inline calculation scheme in CMAQ 48 was applied. In addition to pH changes by mineral dust, the dust-derived Fe and Mn were considered in Exp. B. The solubility of dust Fe and Mn was set as 1 and 50%, respectively 33 (Table 1). The details of the dust inline calculation are given in the Supporting Information (Section S3).
In Exp. C, increased TMI solubilities were considered because one of the uncertainties in modeling settings still under debate is TMI solubilities. Based on a literature review, the range of TMI solubility for Fe and Mn are 0.03−54 and 1.2−97%, respectively. 14 The solubilities of anthropogenic Fe and Mn were set at 54 and 97%, respectively, for the maximum possible production through TMI processes (Table 1). Based on the integrated massively parallel atmospheric chemical transport (IMPACT) global aerosol model 41  (1) In these equations, [SO 4 2− ] i is the SO 4 2− concentration from each process and F i represents the fractional contribution for each process, where i indicates the SO 4 2− formation process.  4 2− produced by AQ(NO 2 ) is expected based on the following three mechanisms: radical chain mechanism, 50 oxygen-atom transfer from OH − , 51 or from O 2 . 52 Additionally, a detailed discussion of the uncertainties of the AQ(O 3 ) end member is given in the Supporting Information (Section S5 and Figure S5). The SO 4 2− derived from the boundary conditions is considered the background existing SO 4 2− , and thus monthly values observed at Alert, Canada, ranging from 0.5 to 1.3‰ 33  ) was statistically evaluated. The metrics were correlation coefficient (R), normalized mean bias (NMB), and normalized mean error (NME).
Here, N is the total number of paired observations (O) and models (M), and these averages are denoted as O̅ and M̅ , respectively. The recommended metrics for SO 4 2− concentration are model performance goals for best performance of R > 0.7, NMB < ±10%, and NME < +35% and model performance criteria for acceptable performance of R > 0.4, NMB < ±30%, and NME < +50%. 53 over East Asia was divided into three seasons, late winter to spring (February−May), summer (June−August), and autumn to winter (September−December), to characterize seasonal behaviors ( Figure 2). The highest SO 4 2− concentrations were found over mainland China, and the higher-concentration regions extended into the downstream region over the Korean Peninsula and Japan. This feature was dominant from spring to summer; therefore, the observations in the downstream region of Japan detected the polluted air mass resulting from the longrange transport over East Asia. In contrast, from autumn to winter when the strong northwesterly wind field by the Asian monsoon was dominated, the SO 4 2− concentration was low and was characterized by clean background conditions. The weekly and monthly average variations of SO 4 2− concentration observed at NOTOGRO also showed higher concentrations from spring to summer and lower concentrations during autumn to winter (Figure 3a). As the background site in Japan, Chichijima island located in the western North Pacific (i.e., south of Tokyo) showed higher concentration in winter and lower concentration in summer. 54 The difference in the seasonal variation of SO 4 2− concentration seen in NOTGRO is caused by the outflow pattern in East Asia, as found in Figure  2. Overall, throughout the year, the statistical analyses showed that all three modeling experiments (Exps. A, B, and C) generally captured the SO 4 2− concentration ( Table 2) ) values ranging from 1.0 to 1.5‰ were higher than those reported in the polluted region in China, where the observations from Beijing varied from 0.1 to 1.6‰ with a mean of 0.9 ± 0.3‰ 13 and the data from Wuhan varied from 0.14 to 1.02‰ with a timeweighted average of 0.53‰. 23 The observed higher concentration in August probably originated from volcanic eruptions in western Japan, which is discussed in Section 3.2.    (Figure 3b, black line). In Exp. A, GAS was the dominant oxidation process, contributing more than 30% of SO 4 2− formation during the period and reaching up to 80% during spring and summer. Following the GAS process, AQ(H 2 O 2 ) was the next most important process, contributing 10−30% of SO 4 2− formation (Figure 3c). In the model, a higher SO 4 2 was suggested in this study, whereas the importance of AQ(O 2 ) was suggested to be the dominant process in volcanic plumes. 55 This difference indicates that the oxidation processes contributing to SO 4 2− formation inside a plume are different from those contributing to SO 2 oxidation after diffusion into the atmosphere. The discussion of this volcanic impact is given in the Supporting Information (Section S6 and Figure S6 Table 2). For the relative contributions of the SO 2 oxidation processes in the model, AQ(O 3 ) increased, whereas GAS decreased ( Figure  3d). This switch can be explained by the neutralization of atmospheric acidity by dust-derived CaCO 3 , and the increased pH obtained by including the inline dust calculation is consistent with previous works. 22 4 2− ), the increases in TMI solubilities in the polluted air were considered in Exp. C. In the standard model of Exp. A, although the anthropogenic emission of TMIs is accurately considered based on the latest emission data set, the modeled estimates of solubilities of anthropogenic and dust Fe and Mn remain highly uncertain in cloud water, as pointed out previously. 14,29,31,33 The increased Fe solubilities have been indicated based on global model comparisons with multiple field campaigns over the Northwest Pacific 57 and also observed toward the downstream region of East Asia, 58,59 mainly because fine mineral aerosols can be acidified due to air pollution. Thus, for Exp. C, we implemented the higher solubilities for both anthropogenic and dust Fe. Compared to the anthropogenic Fe solubility (10%) and Mn solubility (50%) considered in Exps. A and B, the maximum solubility for anthropogenic Fe (54%) and Mn (97%) over the literature 14 was taken in Exp. C ( Table 1). Note that the higher solubility of Fe in dust (3%) considered in Exp. C is generally consistent with that simulated by the IMPACT global aerosol model 57 (Figures S1 and S2 in the Supporting Information).

Disagreement between Observed and Modeled
The SO 4 2− concentration in Exp. C did not increase substantially compared with that in Exps. A and B (Figure 3a and Table 2), but the overestimation in modeled Δ 17 O(SO 4 2− ) from February to March was clearly decreased in Exp. C (Figure 3b, orange line; Table 2) and ranged from 0.54 to 1.94‰. The improvement in Δ 17 O(SO 4 2− ) values was explained by the increased contribution of AQ(O 2 ) owing to the higher solubilities of Fe and Mn (Figure 3e), especially from October to December. Regarding the importance of the AQ(O 2 ) process, TMI concentration levels have been discussed in China. In Beijing, TMI-catalyzed oxidation showed a clear distinguishment of higher/lower contribution during polluted/clean periods. 14 In Wuhan, the enhanced role of TMI-catalyzed oxidation in winter was suggested due to higher PM 2.5 concentration. 23  concentration was higher and from October to December when SO 4 2− concentration was lower. Therefore, it is concluded that the role of TMI-catalyzed AQ(O 2 ) by enhancing solubilities does not depend on the pollution level in this case. Overall, the increased pH obtained by including dust and the increase in TMI solubilities in Exp. C showed the best match for SO 4 2− concentration and Δ 17 O(SO 4 2− ) among three experiments conducted in this study. The series of results strongly indicate the importance of dust and TMI solubilities for SO 4 2− formation via changes in oxidation processes in downstream regions of East Asia, which was not discovered in previous studies that considered only SO 4 2− mass.  Table S2 in the Supporting Information) and Δ 17 O(SO 4 2− ) values were overestimated, even in Exp. C (observed values: 1.24 ± 0.20 and 1.26 ± 0.33‰; modeled values in Exp. C: 1.94 ± 0.09 and 1.57 ± 0.08‰; Table S3 in the Supporting Information). Since the domestic contribution to SO 4 2− in Japan is estimated to be small except in summer, 8 ). In terms of the SO 4 2− formation process, faster H 2 O 2 oxidation of SO 4 2− formation in high solute strength was suggested in Chinese haze events. 21 The current CTMs are based on kinetics research in dilute aqueous solutions and may miss such strengthened features in the atmosphere. The contribution of AQ(H 2 O 2 ) was declined in Exp. C compared to that in Exp. A (Figure 3e,c) through increased dust-derived pH and enhanced TMI solubilities in this study. Because AQ(H 2 O 2 ) does not depend on pH, faster oxidation by H 2 O 2 in Chinese pollution would also affect the downstream region over East Asia even in the application of Exp. C. In this study, although we do not include reactions on the aerosol surface because our focus is not on the Chinese haze itself, the accurate modeling to capture the enhanced SO 4 2− concentration in haze events will also improve SO 4 2− behaviors in the downstream region if the long-range transport occurs.

Toward Closer Agreement between
To date, SO 4 2− production mechanisms have been proposed to explain this missing formation in Chinese pollution and haze events, as introduced. Although this study cannot identify a single mechanism for this missing formation pathway, our results imply that the oxidation pathways for inadequate SO  ) and TMI solubilities along trajectories of long-range-transported air mass from upwind to downstream in East Asia will be one significant approach. These accurately modeled SO 4 2− formation processes over East Asia are necessary to build SO 2 emission regulation strategies along with carbon neutrality 60 because the unknown response to SO 2 emission reduction has already been reported in Western countries. 27,28 The changes in SO 4 2− oxidation processes also alter the size distribution of SO 4 2− and hence direct and indirect radiative forcing, 61,62 closely related to climate aspects. The reduction of SO 4 2− in future atmospheric conditions (i.e., higher CO 2 concentration and lower SO 2 concentration) will increase atmospheric warming compared with the current atmospheric conditions through the slow climate response. 63,64 The reduction of SO 4 2− will also reduce atmospheric acidity and alter the magnitude, distribution, and deposition mode of nutrients supplied to the ocean in the coming decades. 65 As we verified the important role of Fe as the catalyst on the SO 4 2− formation, the declined acidity in the future will relate to weakening the role of TMI-related SO 4 2− formation in the downstream of dust sources over East Asia. Given that a significant emission reduction of CO 2 combined with a welldesigned emission pathway of SO 2 is required, 66 our findings on the role of dust and TMI solubilities and the way to improve the modeling of SO 4 2− formation will contribute to better emission regulations required for air quality and climate change.
■ ASSOCIATED CONTENT

* sı Supporting Information
The Supporting Information is available free of charge at https://pubs.acs.org/doi/10.1021/acs.est.2c03574. Setup of regional chemical transport modeling (Section S1); configuration of CMAQ chemical mechanism for SO 4 2− (Section S2); description of dust inline calculation over East Asia (Section S3); discussion on TMI solubilities (Section S4); uncertainty analysis of